Treatment of highly polluted industrial wastewater by means of sequential aerobic biological oxidation-ozone based AOPs

Chemical Engineering Journal 361 (2019) 89–98

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Treatment of highly polluted industrial wastewater by means of sequential aerobic biological oxidation-ozone based AOPs

T

A.M. Cháveza, O. Gimenoa, , A. Reya, G. Pliegob, A.L. Oropesac,d, P.M. Álvareza, F.J. Beltrána ⁎

a

Departamento de Ingeniería Química y Química Física, Instituto Universitario de Investigación del Agua, Cambio Climático y Sostenibilidad (IACYS), Universidad de Extremadura, Av. de Elvas S/N, 06006 Badajoz, Spain b Sección Departamental de Ingeniería Química, Facultad de Ciencias, Universidad Autónoma de Madrid, Cantoblanco, 28049 Madrid, Spain c Unidad de Toxicología, Departamento de Sanidad Animal, Universidad de Extremadura, Av. de Elvas S/N, 06006 Badajoz, Spain d Instituto Universitario de Investigación en Biotecnología Ganadera y Cinegética (INBIO G+C), Universidad de Extremadura, 10003 Cáceres, Spain

HIGHLIGHTS

GRAPHICAL ABSTRACT

biological-chemical treatment of a • Acomplex industrial wastewater has been studied.

Hazardous wastewater consisted • Real of a mixture of special chemical effluents.

of the aerobic culture • Acclimation used was successfully investigated. photocatalytic ozonation was • Solar able to remove biorecalcitrant compounds.

ARTICLE INFO

ABSTRACT

Keywords: Hazardous wastewater Integrated treatment Biological oxidation Photocatalytic ozonation

The feasibility of the treatment of a complex industrial wastewater by aerobic biodegradation in a sequential batch reactor (SBR) followed by ozone-based advanced oxidation processes (AOPs) has been studied. The industrial wastewater had high organic load (TOC > 3 g L−1, COD > 12 g L−1, BOD5 > 2 g L−1) including some toxic/harmful compounds and high concentration of metal and other inorganic species. SBR treatment of the industrial wastewater diluted with urban wastewater (dilution 1:5), was successful after complete acclimation of the mixed culture (i.e., > 50% COD and TOC removals). Nevertheless, the SBR effluent was still not acceptable to be disposed into the environment (c.a. COD 850 mg L−1) so ozonation, solar photo-ozonation and solar photocatalytic ozonation processes were investigated as further polishing treatments. Thus, the sequential combination of aerobic biodegradation and solar photocatalytic ozonation with a TiO2-based catalyst led to an effluent suitable for discharge into the aquatic environment according to environmental regulations (COD < 125 mg L−1, BOD5 < 25 mg L−1).

1. Introduction The management and treatment of wastewater from residential and industrial sites is critical to the sustainability of water systems. It is

⁎

crucial for treatment technologies to be as much efficient and economically feasible as possible [1]. To deal with hazardous effluents from some industrial sites, such as chemical multi-product plants with diverse toxic wastewater streams, incineration is sometimes considered

Corresponding author. E-mail address: [email protected] (O. Gimeno).

https://doi.org/10.1016/j.cej.2018.12.064 Received 25 September 2018; Received in revised form 10 December 2018; Accepted 12 December 2018 Available online 13 December 2018 1385-8947/ © 2018 Elsevier B.V. All rights reserved.

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as the preferred option [2]. However, depending on the characteristics of the effluent, toxic compounds (e.g., VOCs) and metal salts can be formed or remain during incineration. Moreover, incineration facilities are expensive to build, operate, and maintain, being a major cost factor the equipment needed for the flue gas treatment in order to obey international and local environmental regulations [3]. The high cost associated with incineration urges waste generators to seek other alternatives. In this sense, biological treatment is nowadays considered to be among the best available technologies for wastewater treatment because of its low operating and capital costs. However, conventional biological treatments have long retention times and usually fail to degrade high strength wastewater or bio-refractory compounds [4]. Nevertheless, through acclimation, microbes can undergo physiological transformations, resulting in the selection and multiplication of specialised microorganisms capable to resist toxic substances and/or biodegrade recalcitrant substrates [5]. Even so, biological treatment of hazardous wastewaters usually does not lead to an effluent suitable for permitted direct reuse or discharge and for this reason, it has to be assisted by other non-biological technologies, such as advanced oxidation processes (AOPs). Although, the latter imply higher cost, coupling AOPs with biological treatment could be proposed as an appropriate treatment strategy for recalcitrant wastewaters. For instance, Blanco et al. (2012) could efficiently treat a textile wastewater by combining an aerobic SBR with a Fenton process, reducing the influent COD by about 85% [6]. Among AOPs of heterogeneous photocatalytic oxidation has been proved as an efficient one for the degradation of complex organic contaminants [7,8]. However, it is usually a slow process when treating high organic load wastewaters [9]. To enhance the oxidation rate, proper selection of the photocatalyst and the use of oxidising species as electron scavengers and co-oxidants are key factors. Thus, photocatalytic oxidation in the presence of TiO2-based catalysts and ozone (i.e. photocatalytic ozonation) have been proved beneficial in some instances [10]. However, to the best of our knowledge, no work on the integrated biological – photocatalytic ozonation treatment of industrial wastewater is reported in the literature. In this work, the feasibility of a SBR – solar photocatalytic ozonation treatment at lab scale for the purification of a hazardous wastewater from petrochemical and cosmetic products manufacture has been examined. For the photocatalytic ozonation stage, a magnetically separable TiO2 photocatalyst (TiO2 loaded magnetic activated carbon, TiFeAC) with high activity under solar illumination has been used owing to its success in degrading some wastewater contaminants [11,12].

Table 1 Main characteristics of RIW, FIW and MIW effluents. Parameter

RIW

FIW

MIW

pH Turbidity (NTU) Conductivity (mS cm−1) PO43− (mg L−1) SO42− (mg L−1) Cl− (mg L−1) Bicarbonate alkalinity (g CaCO3 L−1) TSS (mg L−1) VSS (mg L−1) COD (g L−1) BOD5 (g L−1) TOC (g L−1) TPC (mg L−1) A465nm UV254nm (samples diluted 1:10) BOD5/COD

5.8 520 4.44 2.6 585 1385 3.56 674 650 12.2 2.2 3.4 113 0.58 1.24 0.18

5.8 420 4.36 2.5 576 1379 3.36 518 508 10.3 1.8 3.3 113 0.58 1.19 0.17

6.6 105 1.28 99 124 307 0.92 < 0.15 < 0.15 2.1 0.8 0.6 23 0.36 0.25 0.38

Table 2 Metal composition of industrial wastewater before and after treatments. Element

FIW

MIW

BW

O3

O3-solar

O3-solar-cat

Mg (mg L−1) K (mg L−1) Ca (mg L−1) Fe (mg L−1) Zn (mg L−1) Mo (mg L−1) Ni (mg L−1) Cr (µg L−1) Cu (µg L−1) As (µg L−1) Se (µg L−1) Cd (µg L−1) Sn (µg L−1) Hg (µg L−1) Pb (µg L−1)

23.6 61.1 277.0 1.42 2.77 2.56 0.20 75.3 6.6 9.4 11.3 0.48 1.41 0.42 1.33

6.8 52.8 57.4 0.45 0.54 0.51 0.04 15.1 1.3 1.9 2.3 0.01 0.28 0.08 0.27

7.3 69.4 32.7 < 0.05 0.21 0.69 0.04 7.1 1.4 2.1 1.9 n.d. 0.30 n.d. n.d.

6.8 61.1 51.6 < 0.05 0.24 0.56 0.04 9.1 1.7 1.8 1.6 n.d. 0.31 n.d. n.d.

7.9 74.5 67.7 < 0.05 0.23 0.59 0.04 15.7 1.9 2.0 1.8 n.d. 0.28 n.d. n.d.

6.9 69.7 40.6 < 0.05 0.11 0.43 0.02 4.8 1.6 1.0 2.2 n.d. 0.33. n.d. n.d.

n.d.: not detected (below quantification limit).

2.2. Biological treatment The activated sludge sample used as starting inoculum was collected from a municipal wastewater treatment plant (MWWTP) at Badajoz city (Spain), which treats domestic wastewater using a conventional activated sludge process. At the time of sampling, the process at the MWWTP operated at average food to microbial ratio (F/M) of 0.5, hydraulic retention time (HRT) of 7 h and mixed liquor volatile suspended solids concentration (MLVSS) of 1.5 g L−1. After collection, the activated sludge sample was allowed to settle at room temperature for 30 min and the supernatant was removed. The bio-solids were transferred to a cylindrical aerobic batch bioreactor (2 L working volume) provided with temperature, pH and dissolved oxygen concentration (DO) controllers. The bioreactor was operated as a sequencing batch reactor (SBR) with fill, reaction, settle, extract and idle stages [14]. The duration of these phases was 15 min, 22 h, 30 min, 15 min and 60 min, respectively. The system was initially fed with SUW only for two days in order to develop a healthy culture. Then, MIW was stepwise introduced in the feed. Thus, at the beginning of each cycle the bioreactor was fed with an increasing volume of MIW and filled up to 2 L with SUW. At the reaction stage, temperature, pH and DO were controlled at 20 ± 2 °C, 7.5 ± 1.0 and 3.5 ± 0.5 mg/L, respectively. Oxygen uptake rate (OUR), sludge volume index (SVI) and effluent COD were measured and used as indicators of the acclimation process success. Once a healthy culture was adapted to MIW, biodegradation experiments were carried out for 10 h (8 h reaction time) to follow COD, TOC, UV254nm and chroma depletion with time. Thus, at time intervals, samples were withdrawn from the reactor

2. Materials and methods 2.1. Industrial wastewater Raw industrial wastewater (RIW) was provided by a waste management company located in Catalonia (Spain). RIW consisted of a mixture of effluents from some industrial sites (manufacture of petrochemical and cosmetic products), making a high-polluted wastewater. RIW was filtered through filter paper (Filter Lab 1305) to remove suspended solids to some extent. Preliminary experiments showed that the filtered industrial wastewater (FIW) was quite recalcitrant to biodegradation, likely because of the presence of metal and organic compounds, which might be toxic to microbes hence inhibiting biodegradation. To assist biodegradation, a synthetic urban wastewater (SUW) was used as growth substrate. SUW was prepared using peptone and glucose as carbon source [13]. Average chemical oxygen demand (COD), total organic carbon (TOC) and inorganic carbon (IC) of SUW were 172, 85 and 47 mg L−1, respectively. RIW was mixed with SUW (dilution 1:5) to prepare a mixed industrial wastewater (MIW), which was further subjected to biodegradation and chemical oxidation experiments. RIW, FIW and MIW were fully characterised. Analyses were carried out in duplicate with mean values shown in Tables 1 and 2. 90

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for analysis. MLSS and MLVSS were measured in the mixed liquor while COD, TOC and UV254 and A465nm absorbance were analysed in samples after solids removal by centrifugation (3000 rpm, Alresa centrifuge, 200 W) and subsequent filtration through 0.45 μm PVDF membrane filters. Some short-chain carboxylic acids and ecotoxicity towards Daphnia magna and Vibrio fischeri were also analysed on MIW before and after the biological treatment. Biodegradation experiments were carried out five times to check for reproducibility.

and malic acids) and some inorganic anions (phosphate, sulphate and chloride) were analysed with by suppressed ion chromatography (Metrohm 881 Compact IC Pro model provided with an ion suppressor and a conductivity detector). A MetroSep A Supp 5 column (150 mm length, 4 mm diameter) at 45 °C was used as stationary phase. Aqueous Na2CO3 was used as mobile phase with a gradient program from 0.6 to 14.6 mM in 50 min and 10 min equilibration time with a flow rate set at 0.7 mL min−1. Some metallic elements were analysed by ICP-MS with a Perkin Elmer NexION 300 apparatus. COD was measured following the standard dichromate reflux method using Hach-Lange commercial cuvettes and a Hach DR2800 spectrophotometer [16]. BOD5 tests were carried out on 500 mL OxiTop® respirometers inoculated with BOD microbe capsules (Cole-Parmer). TOC and IC were determined with a TOC-VCSH Shimadzu analyser. UV254nm and A465nm were determined with an Evolution 201 spectrophotometer from ThermoSpectronic using 1 cm quartz cells. Total phenolic content (TPC) was evaluated by the Folin-Ciocalteau colorimetric method and expressed as phenol equivalents [17]. Identification of the main organic compounds present in wastewater was carried out by means of gas chromatography-mass spectrometry (GC–MS) system operated in electron impact ionization mode to identify the organic compounds present in the effluents. The analyses were performed by gas chromatography/ion trap mass spectrometry (CP-3800/Saturn 2200, Varian, equipped with an automatic injector CP-8200/SPME, solid-phase microextraction). A 30 m length and 0.25 i.d. capillary column (Factor Four VF-5 ms) was used. The carrier gas (helium) flowrate was set at 1 mL min−1. The SPME was carried out with a fibre cartridge (poly(dimethylsiloxane) red), using adsorption and desorption times of 30 min and 5 min, respectively. The sample injection was conducted at 220 °C. The temperature program used was as follows: (i) 40 °C for 5 min; (ii) then from 40 °C to final temperature of 300 °C at 15 °C min−1; (iii) held at 300 °C for 2 min. Compounds identification was assessed using the National Institute of Standards and Technology (NIST) database. Acute toxicity tests with Vibrio fischeri and Daphnia magna were also carried out. Luminotox® was used to evaluate the bioluminescence inhibition of the marine bacteria V. fischeri (LUMIStox 300, Dr. Lange) following ISO 11348-2 [18]. D magna acute toxicity tests were conducted with the cladoceran D. magna Straus following the protocols given by Baird et al. [19] and the OECD Guideline 202 [20]. Toxicity results were expressed in terms of EC50(%) and TU (Equitox m−3). Values of EC50 (%) were calculated using LUMISsoft 4 SoftwareTM for V. fischeri tests and using a probit analysis (MINITAB STATISTICAL SoftwareTM 2000) for D. magna tests. Mixed liquor samples from the bioreactor were tested for mixed liquor suspended solids (MLSS), mixed liquor volatile suspended solids (MLVSS), sludge volume index (SVI) and specific oxygen uptake rate (sOUR) following standard methods [15]. In ozone-based AOPs, ozone concentration in aqueous solution was determined by the indigo method [21].

2.3. Advanced oxidation processes The biologically treated wastewater (BW) was further subjected to different ozone-based AOPs such as single ozonation (O3), solar photoozonation (O3-solar) and photocatalytic ozonation (O3-solar-cat). All the experiments were carried out in semi-batch mode, using a cylindrical glass-made reactor equipped with a magnetic stirring system, a gas diffuser and gas inlet, gas outlet and liquid sampling ports. A solar box (Suntest CPS, Atlas) provided with a 1500 W Xe lamp and cut-off filters (λ = 300–800 nm, irradiation intensity 550 W m2) was also used. In a typical experiment, the reactor was first loaded with 0.65 L of BW and the required amount of catalyst. If required (experiments in the dark), the reactor was covered with aluminium foil. The reactor content was stirred for 30 min before switching on the Xe lamp and providing a continuous flow (20 L h−1) of an ozone-oxygen mixture (20–30 mg O3 L−1) to the reactor. Ozone concentration at the entrance and exit of the reactor was monitored with in-line ozone analysers (Anseros GM-6000-OEM and GM6000-PRO models, respectively) Also, ozone concentration in solution was measured. Experiments lasted for 5–8 h and samples were withdrawn from the reactor at time intervals to follow parameters such as pH, UV254nm, COD, TOC and phenolic compounds (TPC). The temperature was kept constant at 37 ± 2 °C throughout the experiments. To assess the impact of carbonate/bicarbonate ions present in BW on the efficiency of ozone-based AOPs, some experiments were carried out in the absence of inorganic carbon. To do so, prior to the degradation experiments carbonate/bicarbonate ions were removed from BW by means of air-striping at acid pH. Once the striping operation was completed, the pH was restored to circumneutral conditions. The catalyst used in photocatalytic experiments was a TiO2-Fe3O4activated carbon magnetic composite (TiFeAC) synthesized according to a previous work [12]. Briefly, magnetic carbon particles were obtained by impregnation of a commercial activated carbon (Darco 12–20, Sigma Aldrich) with an iron (III) nitrate ethanol solution followed by iron reduction with ethylene glycol and heat treatment at 550 °C for 4 h. Separately, a TiO2 nanosol was produced from titanium (IV) butoxide. Finally, the magnetic carbon particles were dispersed in the TiO2 nanosol under sonication for 1 h. The product was dried under vacuum at 80 °C and washed thoroughly with distilled water to remove impurities. TiFeAC composition was 70.0 wt% TiO2 (anatase), 7.5 wt% Fe (mainly as magnetite and maghemite) and 18.5 wt% activated carbon. This catalyst was chosen because of its good photocatalytic activity, stability and facile recovery using a magnet [12]. To assess the stability and reusability of the photocatalyst, a series of five consecutive solar ozonation experiments was carried out. After each run, the catalyst was separated with a magnet, washed with distilled water and used to treat fresh BW in the next run. TOC, COD, UV254nm and Fe were measured in the treated effluent.

3. Results and discussion 3.1. Industrial wastewater characterisation Table 1 shows the main characteristics of RIW, FIW and MIW samples used in this study. RIW could be classified as a high strength effluent in terms of organic load because of high COD, TOC and TPC values. In addition, the industrial wastewater showed high conductivity mainly as a result of the presence of large concentrations of bicarbonate, sulphate, chloride and phosphate ions. The biodegradability index, measured as the BOD5/COD ratio, was below 0.2, which means poor biodegradability [22]. Upon filtration of RIW, suspended solids and particulate organic matter were removed to some extent (note lower values of TSS, COD, BOD5 and TOC in FIW in comparison to RIW). However, the concentrations of inorganic ions, conductivity and the biodegradability index remained practically unchanged from RIW

2.4. Analytical methods Turbidity, pH, conductivity and dissolved oxygen were measured with a Hanna HI 93,414 turbidity-meter, a Crison GLP21 + pH-meter, a 524 Crison conductivity-meter and an HQd Portable Hach meter, respectively. Total and volatile suspended solids (TSS and VSS, respectively) were measured gravimetrically following standard methods [15]. Short-chain organic acids (oxalic, acetic, pyruvic, formic, succinic 91

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acclimation of microorganisms to the industrial effluent. The MLVSS/ MLSS ratio was kept within a typical range of 0.75–0.9 [25]. Table 5 shows the SBR process performance during the acclimation stage. As seen in Table 5, when the activated sludge sample was brought into contact with FIW for the first time (day 1) an abrupt decrease in the aerobic biological activity was observed as sOUR fell from 52 to 9.1 mg O2 g VSS−1 h−1. OUR inhibition was calculated as follows [26]:

1.0

COD/COD0

0.8 0.6 0.4 0.2 0.0

Inhibition (%) = 1

Cycle 1 Cycle 2 Cycle 3 Cycle 4 Cycle 5

0

1

2

3

4

5

6

7

sOUR × 100 sOURb

(1)

where sOUR and sOURb are referred to the sample exposed to industrial wastewater and the sample exposed to SUW (used as blank wastewater). sOURb was in the 47–55 mg O2 g VSS−1 h−1 range. For 12.5% FIW influent wastewater, high inhibition values (82.5% and 80.6%, respectively) were recorded in days 1 and 2 of acclimation. However, inhibition was drastically reduced from day 2 to 3 (48.8%), indicating a better performance of the activated sludge process. As a general rule, for a given percentage of MIW in the influent wastewater, increasing the time of acclimation resulted in higher sOUR (decreasing inhibition). At the end of the acclimation stage (day 15, 100% MIW), relatively low OUR inhibition (c.a. 30%) was found. Process performance during the acclimation stage was also followed by COD removal. As it is apparent in Table 5, the percentage of COD removal was relatively low at the beginning of the acclimation phase (< 40% removal in days 1 and 2) but increased up to about 60% with the acclimation time. An adapted population was consistent with the effluent COD concentration reaching constant values and the MLVSS concentration gradually increasing [27]. SVI was used to characterise the settling properties of the activated sludge during the acclimation stage. In practice, SVI can vary from 30 to 400 mL g−1 [28]. As a rule, a proper SVI value, typically below 100 mL g−1, is an indicator of good settling properties of the sludge [29]. Therefore, keeping SVI below 100 mL g−1 is crucial for the optimal operation of an activated sludge process. As seen in Table 5, SVI values obtained during the acclimation phase varied from 163 to 54 mL g−1 from day 1 to 15. The presence of metals and hazardous compounds resulted in poor sludge settling behaviour from day 1 to 5 (SVI > 150 mL g−1). However, SVI was kept below 100 mL g−1 from day 10 onwards. This also reveals the success of the sludge acclimation to MIW.

8

Time (h) Fig. 1. Total ion current GC–MS chromatogram of (a) FIW, (b) MIW and (c) BW. For main wastewater characteristics see Table 1.

to FIW. Low biodegradability of the wastewater can be attributed to the high organic load and conductivity and to the presence of toxic organic compounds such as TPC. In addition, noticeably concentration of some heavy metals was measured in FIW as shown in Table 2, being Fe (1.42 mg L−1), Zn (2.77 mg L−1) and Mo (2.56 mg L−1) the most abundant ones. UV254nm absorbance of RIW and FIW was high (note that UV254nm values shown in Table 1 were obtained on diluted samples), which suggests the presence of aromatic and unsaturated compounds in the effluent. In fact, some of these chemical structures were detected in FIW by GC–MS as shown in Fig. 1 and Table 3. The GC–MS results reveal the presence of a number of compounds related to petrochemical and cosmetic industries, which agrees with the origin of the industrial wastewater. Table 4 shows results of the industrial wastewater toxicity towards Vibrio fisheri and Daphnia magna. Low EC50 and high TU values were obtained in the toxicity tests of FIW, indicating that the industrial wastewater was toxic to both organisms [23]. The characteristics of the industrial wastewater discussed above make it unsuitable for discharge into the municipal sewer system. According to local regulations, maximum allowable limits for discharge are established at: COD = 1.5 g L−1; BOD5 = 0.75 g L−1; TPC = 2 mg L−1; Toxicity towards V. fischeri = 25 Equitox m−3 among other parameters [24]. Therefore, wastewater treatment is compulsory. Given the low wastewater biodegradability, direct biological treatment cannot be recommended. As a cost-effective treatment, dilution of the industrial wastewater with urban wastewater and a sequential biological – chemical treatment of the combined effluent is proposed in this work in order to fulfil environmental regulations for discharge into the aquatic environment (pH = 6–8; COD = 125 mg L−1; BOD5 = 25 mg L−1). Dilution of FIW with SUW resulted in the MIW effluent, which presented reduced COD, BOD5 and TPC values compared to FIW but still did not meet limit regulations for discharge. However, it is important to note that the biodegradability index (i.e., BOD5/COD) of the industrial wastewater increased as a result of dilution (see Table 1), thus making the MIW effluent more amenable to biological degradation. Also MIW toxicity towards V. fisheri and D. magna was much lower than that of FIW (see Table 2).

3.2.2. MIW biodegradation Additional MIW biodegradation experiments were carried out in the SBR using acclimated activated sludge. Fig. 2 shows the evolution of residual COD during the biodegradation stage of five consecutive SBR cycles (reaction time 8 h) where the biomass concentration was adjusted at the beginning of each cycle to about 1.8 g VSS L−1. The gradual decline of COD indicates good degrading activity of the acclimated sludge. The evolution of TOC also followed a similar trend (results not shown). On average, COD and TOC removals after the 8-hour biodegradation stage were about 50 ± 5% and 53 ± 4%, respectively. Chroma removal, which was determined as the percentage of A465nm disappearance upon the treatment, was about 90%, suggesting the biodegradation of colouring substances. Regarding metals, only Fe and Zn have been removed to some extent after the biotreatment, probably due to adsorption onto activated sludge [30]. GC–MS analysis of biotreated samples (BW) showed that only three out of twenty compounds detected in MIW were present in BW. As shown in Table 3, the compounds that were not biodegraded by the acclimated culture were C-88 (metyl-tert-butyl-ether), C-9 (m-methyl pyridine) and C-262 (2,6-bis (1,1-dimethylethyl)-4-(1-oxopropyl)phenol). Likely as a result of this, toxicity of BW towards V. fisheri was reduced significantly with regard to MIW. However, little changes were observed in the ecotoxicity tests conducted with D. magna on MIW and BW (Table 3).

3.2. Aerobic biological oxidation 3.2.1. Activated sludge acclimation A 15-day acclimation phase was considered in this study to adapt the microorganisms of the activated sludge sample taken from the MWWTP to the industrial wastewater. MIW was included in the feed of the activated sludge system from day 1 and its concentration was gradually increased throughout the acclimation phase, revealing the 92

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Table 3 Main organic compounds in FIW, MIW and BW as detected by GC–MS. Compound

tR (min)

Molecular formula

Molecular weight

Peak 1 C-88

1.674

C5H12O

88

Peak 2 C-102

2.313

C5H10O2

102

Peak 3 C-116

3.115

C6H12O2

116

Peak 4 C-100

3.487

C6H12O

100

Peak 5 C-92

4.117

C7H8

92

Peak 6 C-160

4.743

C9H20O2

160

Peak 7 C-116

5.385

C6H12O2

116

Peak 8 C-93

6.154

C6H7N

93

Peak 9 C-120

8.101

C9H12

120

Peak 10 C-130

8.577

C8H18O

130

Peak 11 C-152

9.091

C8H8O3

152

Peak 12 C-142

11.711

C11H10

142

Peak 13 C-142

11.874

C11H10

142

Peak 14 C-166

12.440

C9H10O3

166

Peak 15 C-154

12.526

C12H10

154

Peak 16 C-170

12.717

C12H10O

170

Tentative structure

Samples(1)

Methyl-tert-butyl-ether

FIW MIW BW

Propyl acetate

Methyl trimethylacetate

Methyl isobutyl ketone

Toluene

2,2'[Methylenebis(oxy)]bis(2methylpropane)

Butyl acetate

m-methyl pyridine (3-picoline, 3 Mepy)

1,2,4-trimethylbenzene

2-ethyl-1-hexanol

2,2-dihydroxy-1-phenylethanone

1-methyl-naphthalene

2-methyl-naphthalene

3-methoxy-benzoic acid methyl ester

Biphenyl

FIW MIW FIW MIW

FIW MIW FIW MIW FIW MIW

FIW MIW FIW MIW BW FIW MIW

FIW MIW

FIW MIW

FIW MIW

FIW MIW FIW MIW

FIW MIW FIW MIW

Diphenyl oxide

(continued on next page)

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Table 3 (continued) Compound

tR (min)

Molecular formula

Molecular weight

Peak 17 C-192

13.051

C12H16O2

192

Peak 18(2) C-226

14.278

C16H18O

226

Peak 19(2) C-226

14.450

C16H18O

226

Peak 20 C-262

14.543

C17H26O2

262

Samples(1)

Tentative structure

FIW MIW

Methyl 4-tertbutylbenzoate

FIW MIW

Bis(1-phenylethyl) ether

Bis(1-phenylethyl) ether

FIW MIW FIW MIW BW

2,6-Bis(1,1dimethylethyl)-4-(1oxopropyl)phenol (1) (2)

Samples where compounds were detected. Isomers.

solar) and solar photocatalytic ozonation (O3-solar-cat). A comparison of process performance in terms of COD, TOC and TPC removals is presented in Fig. 3. The highest COD and TOC depletion rates were observed in photocatalytic ozonation. Single ozonation gave rise to a poor mineralization degree (TOC removal < 15%) and COD removal (< 25%) after 5 h reaction time. On the other hand, photo-ozonation (O3-solar) and, especially, photocatalytic ozonation led to significant TOC and COD removal percentages (25–45% and 40–50%, respectively). This can be attributed to the increased generation of secondary oxidising species (e.g., hydroxyl free radicals) in the photo-treatments compared to single ozonation. In fact, dissolved ozone observed during O3-solar and O3-solar-cat treatments was around 5 μM compared to 30 μM reached in single ozonation runs, which suggests the efficient photo-decomposition of O3 into HO· radicals [31]. The photocatalytic treatment (O3-solar-cat) showed a superior power in terms of TOC and COD removals, due to the promotion of different catalytic pathways of HO· radicals production as shown in a previous work [12]. The evolution of TPC (Fig. 3(c)) was similar in all the treatments with nearly 90% TPC removal after 5 h of treatment. In connection with this, UV254nm was also removed to a great extent (> 90%) and complete elimination of the organic compounds detected by GC–MS in BW was achieved with

Table 4 Ecotoxicity results of Vibrio fischeri inhibition and Daphnia magna immobilization tests after exposure to treated and untreated industrial wastewater. Sample

FIW MIW BW O3 O3-solar O3-solar-cat

Vibrio fischeri

Daphnia magna −3

EC50 (%)

TU (Equitox m

1.6 9.2 17.1 13.3 5.7 12.3

62.5 10.9 5.8 7.5 17.5 8.1

)

EC50 (%)

TU (Equitox m−3)

3.2 25.3 24.0 46.9 50.5 74.8

31.2 3.9 4.2 2.1 2.0 1.3

3.3. Advanced oxidation processes Although the biotreatment of MIW was satisfactory to some extent, BW presented pollution in terms of TPC (15 mg L−1 on average) and, COD (850 mg L−1 on average) and BOD5 (100 mg L−1) well above the allowable discharge limit into the aquatic environment (Table 6). In order to further remove COD, TOC and TPC from BW, various ozone AOPs were tested: single ozonation (O3), solar photo-ozonation (O3Table 5 Activated sludge process performance during the acclimation stage. Day

% MIW in the influent wastewater

MLSS (mg L−1)

MLVSS (mg L−1)

sOUR (mgO2 gVSS−1 h−1)

% Inhibition

% COD8h removal

% COD22h removal

SVI (mL g−1)

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15

0 12.5 12.5 12.5 25 25 37.5 37.5 50 50 75 75 100 100 100 100

1093 1107 1115 1100 1087 1078 1247 1057 1207 1123 1327 1330 1678 1910 2098 2033

872 928 898 847 855 862 962 932 1065 955 1148 1165 1478 1787 1785 1780

52.0 9.1 10.1 26.6 36.7 38.5 38.4 39.8 40.7 43.5 32.6 40.0 25.4 30.0 35.8 35.9

0 82.5 80.6 48.8 29.2 25.9 26.1 23.4 21.6 16.3 37.1 22.9 51.2 42.1 31.1 30.8

– – – – – – – – 46,0 44,6 51,2 49,1 46,6 48,4 56,1 57,0

30.1 26.0 35.4 63.1 44.9 59.2 49.6 54.6 56.2 53.8 61.1 62.8 59.9 64.3 61.9 61.3

100 163 161 164 146 158 104 123 108 116 98 68 60 58 55 54

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Fig. 2. Evolution of normalized residual COD during biodegradation of MIW in a SBR. Experimental conditions: pH = 7; T = 20 °C; Initial MLVSS = 1.8 g L−1. For main MIW characteristics see Table 1.

beneficial impact on the photo-ozonation process performance in terms of COD and TOC removals. As seen in Fig. 4(c), where the evolution of pH throughout the experiments is presented, the removal of IC from BW avoided the buffering effect observed in the as-obtained BW. In fact, pH values were in the range 7.5–8 during the entire reaction time in experiments completed with as-obtained BW in contrast to those with ICfree BW, where pH varied from 7 to 4.3. Therefore, despite the hydroxyl radical scavenging effect of bicarbonate, it seems that neutral-basic pH, which favours indirect ozonation reactions with organics, plays a more important role than bicarbonate itself [33]. TOC was about three times higher than IC in BW. Then, assuming an average rate constant for any organic compound in the range kHO = 107–109 M−1s−1, the apparent reaction rate of HO% radicals with the organic compounds is expected to be higher than that of the bicarbonate-HO% reaction. Moreover, the rate of some direct ozone-organic compounds reactions is favoured with increasing pH, which also contributes to BW degradation [34,35]. Additionally, a series of five consecutive runs was carried out reusing the TiFeAC photocatalyst. Fig. 5 shows the TOC and COD removal percentages after the reusability runs. As it is apparent, the photocatalyst kept its activity throughout the entire series of runs (67% and 74% TOC and COD removals on average). Furthermore, the concentration of iron found in solution was below 0.05 mg L−1 which demonstrates the stability of the photocatalyst under the reaction conditions. Finally, to improve the COD and TOC removals achieved by solar photocatalytic ozonation, additional experiments were carried out with higher loading of the photocatalyst and higher ozone dose. Results in terms of COD and TOC removal are depicted in Fig. 6(a) and (b), respectively. A slight increase in process performance was observed when the catalyst loading was two-folded, achieving a higher COD conversion (74% vs. 64%) though no clear improvement was observed in TOC removal. Regarding the ozone dose, an increase in the ozone applied from 0.6 to 0.9 g L−1 h−1 led to an enhancement in both the COD (from 64 to 84%) and TOC (from 60 to 70%) removals As a result of the photocatalytic treatment a number of short-chain carboxylic acids were generated, which accounted for 30%, 40% and 48% of overall TOC in

Table 6 Main characteristic of BW and the O3-solar-cat effluent and maximum allowable limits for effluent discharge into the municipal sewer system (A) and into the aquatic environment (B). Parameter

BW

O3-solarcat

Maximum allowable limit A

Maximum allowable limit B

pH COD (mg L−1) BOD5 (mg L−1) TOC (mg L−1) TPC (mg L−1) UV254nm (dil. 1:10)

7.2 850 100 280 15 0.113

6.7 118 < 20 67 0.7 0.013

6–10 1500 750 – 2 –

6–8 125 25 – – –

all the ozonation treatments. Some short-chain organic acids were identified in ozone-treated samples. Oxalic, pyruvic, acetic and formic acids accounted for around 10%, 25% and 30% of the overall TOC remaining after the O3, O3-solar and O3-solar-cat processes, respectively, being major by-products of the treatments. In addition to COD, TOC and TPC, chroma was also removed to a high extent by the AOPs applied. Thus, chroma removal efficiencies achieved in the sequential SBR-AOP processes were higher than 99% regardless of ozone-based AOP considered. The toxicity of the final effluent was also evaluated and results are shown in Table 4. For V. fischeri, EC50 values did not follow a clear trend in any case. However, for D. magna a clear decrease of the toxicity was observed when the AOPs were applied (especially for O3-solar-cat), indicating the efficiency of the sequential biologicalchemical treatment in the detoxification of MIW. It is well known that carbonate or bicarbonate ions may act as HO· scavengers in AOPs (kCO32 HO = 3.7 × 108 M−1s−1; kHCO3 HO = 2 × 107 M−1s−1 [32]). Since IC was high in BW (90 mg L−1 on average), the effect of carbonate/bicarbonate ions removal was studied to analyse the impact of these species on the efficiency of the photo-ozonation and photocatalytic ozonation treatments. Fig. 4 compares COD and TOC removals and pH evolution during experiments carried out with asobtained BW and IC-free BW. From Fig. 4(a) and (b) it is apparent that, contrary to what was expected, the presence of bicarbonate in BW had a 95

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Fig. 4. Effect of inorganic carbon on the evolution of normalized COD (a), TOC (b) and pH (c) during ozone-solar radiation based AOPs. Experimental conditions: T = 37 °C; V = 0.65 L; Qg = 20 L h−1; CO3,g = 20 mg L−1; I = 550 W m−2 (λ = 300–800 nm); Ccat = 0.375 g L−1 (if catalyst applied). For main BW characteristics see Table 6.

Fig. 3. Evolution of normalized residual COD (a), TOC (b) and TPC (c) during the application of various AOPs to treat BW. Experimental conditions: T = 37 °C; V = 0.65 L; Qg = 20 L h−1; CO3,g = 20 mg L−1; I = 550 W m−2 (λ = 300–800 nm) (if radiation applied); Ccat = 0.375 g L−1 (if catalyst applied). For main BW characteristics see Table 6.

Table 6. It can be seen that the effluent after the sequential treatment fulfils legal conditions for direct discharge into the environment in terms of pH, COD and BOD. Also, TOC, TPC and UV254nm were satisfactory. In addition, no important concentration of heavy metals was present in the treated effluent (see Table 2) and all the organic compounds detected by GC–MS in FIW were completely removed (see Table 3). Blanco et al. (2012) also found successful an integrated aerobic SBR-

the final effluent after 8 h of treatment at the conditions (a) 20 mg L−1 O3, 0.375 g L−1 catalyst; (b) 20 mg L−1 O3, 0.750 g L−1 catalyst; and (c) 30 mg L−1 O3, 0.375 g L−1 catalyst, respectively. At the most stringent oxidation conditions of those studied here (O3solar-cat, 30 mg L−1 O3, 0.375 g L−1 catalyst) final values of the main characteristic parameters of treated wastewater are summarized in

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100

TOC COD

90 80 Removal (%)

70 60 50 40 30 20 10 0

1

2

3

4

5

Run Fig. 5. Photocatalyst reusability in O3-solar-cat treatment. Experimental conditions: t = 5 h; T = 37 °C; Qg = 20 L h−1; Ozone dose (each run) = 4 g L−1h−1; I = 550 W m−2 (λ = 300–800 nm); Ccat = 0.375 g L−1. For main BW characteristics see Table 6.

Fenton process to treat a textile industrial wastewater, with similar organic load than that considered in this work. They also achieved high overall COD and TOC removals, averaging 86% and 92%, respectively, which are similar percentages than those found in this work (94% and 89%, respectively at conditions of Table 6) [6]. In the same line than this work, Gimeno et al. (2016) applied an aerobic SBR-photocatalytic ozonation treatment (commercial TiO2/P25 as photocatalyst) to a primary wastewater effluent with spiked pharmaceutical compounds in a pilot plant. They observed about 80% and 70% COD and TOC removals in the overall process leading to a suitable effluent for discharge [36].

Fig. 6. Effect of ozone concentration and catalyst loading on the evolution of normalized residual COD (a) and TOC (b) during photocatalytic ozonation runs to treat BW. Experimental conditions: T = 37 °C; V = 0.65 L; Qg = 20 L h−1; CO3,g = 20–30 mg L−1; I = 550 W m−2 (λ = 300–800 nm); Ccat = 0.375– 0.750 g L−1. For main BW characteristics see Table 6.

4. Conclusions A sequential combination of aerobic biodegradation (SBR) followed by ozone-based AOPs has been successfully used to degrade a highpolluted wastewater from petrochemical and cosmetic industries. The SBR treatment could be successfully applied after dilution of the industrial wastewater with urban wastewater (MIW wastewater). Besides, an acclimation stage was needed to develop an appropriate mixed culture able to biodegrade the components of the industrial wastewater. Thus, while biodegradation of MIW was highly inhibited when using non-acclimated activated sludge (> 80% inhibition), after 15 days of acclimation the activated sludge was able to biodegrade the effluent to high extent. Ozone in combination with solar radiation and a TiO2based photocatalyst (TiFeAC) could further remove recalcitrant compounds remaining after biodegradation as well as to considerably reduce the toxicity of the wastewater, thus obtaining an effluent suitable for direct discharge into the environment. In addition, the photocatalyst was quite stable and could be reused. Therefore, the combination of biological oxidation using acclimated sludge with solar photocatalytic ozonation AOP is a promising approach to decontaminate the hazardous wastewater studied in this work. Further studies at larger scale about the environmental impact and cost estimation of the integrated process would be necessary to definitively establish its feasibility vs incineration, which is the treatment method currently applied at the industrial site.

Acknowledgements This work has been supported by the Spanish Ministerio de Economía y Competitividad (MINECO) and European Feder Funds through the projects CTQ2012-35789-C02-01/PPQ and CTQ201564944-R. Ms Chávez is also thankful to the Spanish MINECO for her predoctoral contract (call 2013). Also, authors wish to thanks the financial support given to Oropesa A.L. by Santander Universidades (Programa 2016-2017: Becas Iberoamérica) and the technical assistance provide by Palma P. and Floro A.M. from Departamento de Tecnologias e Ciências Aplicadas. Escola Superior Agrária de Beja, Beja, Portugal. References [1] I. Alyaseri, J. Zhou, Towards better environmental performance of wastewater sludge treatment using endpoint approach in LCA methodology, Heliyon 3 (2017) e00268, , https://doi.org/10.1016/J.HELIYON.2017.E00268. [2] C.R. Dempsey, E.T. Oppelt, Incineration of hazardous waste: a critical review update, Air Waste 43 (1993) 25–73, https://doi.org/10.1080/1073161X.1993. 10467116. [3] A. Poggio, E. Grieco, Influence of flue gas cleaning system on the energetic

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